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Abstract. Modem large-scale conventional agriculture with intensive monoculture and row-cropping practices often results in unacceptable soil erosion and runoff and in associated losses of nutrients and pesticides It also adversely affects wildlife. Sediment from erosion is the greatest pollutant of surface water in the United States and is a major carrier of agrichemicals in to the water system. Conservation tillage practices can significantly reduce soil losses in modem production systems, but pollution of surface and groundwaters from runoff and associated potential for increased use of pesticides may still present a hazard. Erosion can be reduced to tolerable levers by the use of crop rotations, meadow crops, and mulch tillage, as commonly used in alternative agriculture. These cultural practices also result in diversity in crop types and in smaller fields, with possible benefit to many wildlife species. Chemical threats to the environment and to wildlife are reduced because synthetic chemicals are used sparingly or not at all
The major environmental concerns facing U. S. agriculture today are wind and water erosion of soils and the escape of pesticides and fertilizers into nontarget areas. Of these, soil erosion is by for the most serious and pervasive environmental problem. Soil erosion has been described recently within the U.S. Department of Agriculture as "epidemic in proportion" and a major threat to the nation's soil and water resource base (Brown and Wolf, 1984). Agriculture has now been identified as the leading contributor of nonpoint-source pollution to the nation's waters (Duda and Johnson, 1985). Much of the environmental damage caused by agriculture is irreversible and results in a multibillion dollar drain on the nation's economy. It has been estimated that sediment and associated pollutants, mainly from agricultural lands, cause more than $6 billion in off-site damage each year in the United States (Clark, 1985). This estimate does not include damage to fish and wildlife.
Much of the available information about soil erosion rates in the United States was obtained in the 1970's and is compiled in the National Resources Inventory (NRI) for the year 1977 (USDA, 1981). The NRI was designed to identify soil and water conservation needs. Wind erosion and water erosion from croplands in 1977 were estimated at 0.8 billion metric tons (mt) and 1.7 billion mt, respectively. This total of 2.5 billion mt of erosion for 167 million ha of cropland averages about 15 mt of soil per ha. This value considerably exceeds the tolerance limit (T-value) of 11 mt/ ha that the Soil Conservation Service has judged to be the maximum allowable for maintaining the long-term productivity of the land. Nationally? water erosion alone exceeded the T-value on more than 45 million ha, or 27 percent of the cropland (Larson, 1981).
The NRI survey also showed that erosion was concentrated more in certain areas than in others. For example, water erosion was double the T-value on 9 percent of the land used for crops in the Northwest, 19 percent in the Corn Belt, and 32 percent of the lands used for two crops in the South (Larson, 1981).
The NRI soil loss value is the amount of soil eroded per unit area at a given sample point, but only part of this eroded soil actually leaves the field. For example, a portion of the soil eroded by water is deposited at the base of slopes or in areas with vegetative cover. Movement of soil from one part of the field to another usually results in a change in productivity, but may not cause an adverse environmental effect. An important parameter for off-site environmental impacts is the deposition factor, better known as the delivery ratio; this is the amount (or percentage) of eroded material (sediment in this instance) delivered to waterways. The delivery ratio, although highly variable, is estimated to range between 10 and 40 percent for most agricultural watersheds.
Crosson and Stout (1983) said that a model developed at Resources For the Future predicted that only 40 percent of the soil eroded by water in the United States in 1977 ended up in the nation's waterways. In Iowa, Piest et al. (1977) studied a 30-ha watershed on which the average annual erosion rate was 78 mt/ ha of soil, but only 24 percent of this (19 mt/ha) was actually lost from the watershed. The bulk of the eroded soil was simply moved from higher to lower places in the cultivated fields, where it may have been less valuable for crop productivity.
The same principles apply to wind erosion, in which a high proportion of airborne sediments is deposited on other cultivated lands and, hence, the soil is displaced but not lost for agriculture. However, airborne dust can cover very large areas during storms, and its damage to the environment is much greater than that associated with simple pickup and deposition from field to field.
Although delivery ratios show that much of the eroded soil remains on cultivated lands, the amounts lost from fields are nevertheless substantial. Moreover, the type of material usually carried from the field (mainly soil minerals containing plant nutrients and organic matter) is often more important in regard to the environment than the absolute amount delivered because fine sediments and organic matter carry nutrients and pesticides that can contaminate receiving waters (Frye et al., 1985; Chesters and Schierow, 1985; Wauchope, 1978). These materials also represent substantial losses in soil productivity as they may move from places where they are needed to maintain the depth of the soil mantle to areas where they become a detriment (Frye et al., 1985).
Erosion damage to the environment is mainly concerned with off-site effects, except for aesthetic value degradation associated with rill or gully formation and plant damage by soil deposition. Deposition in fields also harms wildlife habitat by covering nesting areas and wildlife food sources. When soil is carried from fields by runoff or wind, its redeposition occurs mostly at places where the energy of the transport medium (water or air) is reduced so that a portion of the sediment drops out. This will be in areas involving fences, ditches, low-lying areas, ponds, streams, and other surface waters. Sediment in water and resultant siltation not only reduce the efficiency of irrigation, power generation, and navigable systems, but also adversely affect many desirable ecosystems and the use of water for industrial and recreational purposes (Clark, 1985; Duda and Johnson, 1985).
Soil erosion is also a selective or an enrichment process (Stoltenberg and White, 1953). The eroded fraction generally is higher than the original soil in organic matter and clays, where many nutrients, pesticides, and pesticide degradation products are adsorbed or chemically combined (Wauchope, 1978; Foster et al., 1985). Moreover, because of their colloidal nature, organic matter and clays are most likely to remain in runoff water and be carried off the land. The importance of agricultural chemicals in runoff is shown by a recent EPA study of the Chesapeake Bay that concluded, "The nonpoint source runoff from cropland constitutes the largest share of the nonpoint source nutrient load to the Bay" (USEPA, 1983). The study showed that nonpoint sources contribute almost 17 kg/ha/yr of N from the watershed area surrounding the Bay and that about half of this comes from fertilizer N.
Erosion from a specific area can accelerate with time as soil organic matter, sediments, and associated nutrients are lost, resulting in a degradation of soil physical, chemical, and biological properties (Frye et al., 1985). In turn, plant growth is adversely affected because of reduced water intake and storage, reduced nutrient levels, and a poorer rooting medium (Papendick and Elliott, 1984a). Soil productivity must then be maintained with increased amounts of fertilizer to replace the nutrients lost through erosion. If fertilization is done improperly, nutrient levels in eroded materials and runoff may be increased, further increasing the potential for environmental damage.
It is well documented that fertilizers applied to crops are a potential source of nutrient enrichment to surface waters (Myers et al., 1985). Nitrogen is also a potential contaminant of groundwaters. Fertilizer recovery by crops with conventional farming systems is generally far from complete, with about 50 percent recovery of N (Power, 1983). Crop uptake of fertilizer P per season varies between 5 and 25 percent and for fertilizer K from 40 to 75 percent of the amount applied, although more is taken up by subsequent crops (Hauck and Koshino 1971). Nitrogen can be highly mobile it can be tied up in organic master, carried through surface or subsurface water movement, or denitrified. Phosphorous and K, however, attach to soil particles and are carried away only by erosion Langdale et al. (1985) found that losses in runoff from a conventional tillage decreased from about 4 to 0.1 k ha/yr as conservation and cropping practices increased.
Nitrogen fertilizers can contribute NO3 to groundwater reservoirs. This h been documented in Nebraska (Schepers and Mielke, 1983), where concentration of NO3-N in groundwater beneath the Platte River Valley appear to be increasing, and in other states as well (CAST 1985). Concentrations of NO3-N in excess of 10 mg/1 (the upper limit established for drinking water by the U. Environmental Protection Agency (1976)) are commonplace in much Nebraska's irrigated region. Most susceptible to pollution are intensively irrigated corn production areas with shallow water tables and sandy soils, combination that encourages rapid infiltration of NO3-N. However, deep movement of fertilizer N has been documented in nonirrigated soils as well (Schuman et al., 1975).
Other disturbing reports are appearing. In 1984, 40 public water supplies Iowa were put on notice for exceeding the drinking water standard of 10 mg/l as NO3-N (Hallberg, 1984). Included in the list was Des Moines, the large municipality in the United States to exceed this standard to date. Fertilizer applications in the Big Spring Basin in Iowa increased about 250 percent from the 1960's to 1980, while N application as livestock manure were estimated to increase about 30 percent. During the same period, NO3 concentrations groundwater increased by about 230 per cent. There were no industries, landfills municipalities, or large feedlot operations in the Big Spring Basin to co~ plicate the interpretation (Hallberg 1984).
The amounts of NO3-N discharged into groundwater and surface water from the Big Spring Basin for the years 1982 and 1983 were 0.8 and 1.3 million kg for a near-normal and a wet year, respectively. This was equivalent to a 55 and 82.5 kg/ha loss of N. respectively, from the long-term average area in corn in the basin (Hallberg, 1984). These numbers do not imply that all of the N in the runoff and groundwater came from fertilizer N. However, the increase of NO3-N did parallel the increase in fertilizer usage. Although the health hazard of NO3-N in groundwater is not clear (CAST, 1985), the economic value of NO3-N and other forms of N must be considered because it represents a loss to the producer and eventually becomes a cost to society.
Pesticides and pesticide residues can be transported by erosion and runoff to off-site areas in ways similar to those operative in the case of nutrients. These losses are determined by the persistence of the material, the strength with which it is adsorbed to sediments, and its solubility in water (Wauchope, 1978). Although data on a national scale are quite limited, reported pesticide concentrations in streams and lakes are generally very low. The delivery ratio for pesticides to surface waters is generally less than 5 percent (Chesters and Schierow, 1985). Nevertheless, fish kills can occur after intense runoff from croplands where certain highly soluble compounds with low adsorptive capacity have been recently applied (Chesters and Schierow, 1985). However, these authors state that the more soluble pesticides tend to degrade readily to less toxic products, a process which reduces their toxic effects in aquatic systems.
The appearance of pesticides or pesticide residues in groundwaters is a matter of increasing concern. Cohen et al. (1984) listed 12 pesticides or residues found in wells from 19 states. Few have been found at levers above the suggested health advisory concentration, which has a presumed safety factor of 100 (CAST, 1985). Some of the groundwater contamination by pesticides or their residues has resulted from chemicals already removed from the. market. Evidence that some chemicals may leach has caused changes in application practices and rates. CAST (1985) reported that "Analytical techniques are capable of detecting pesticides at concentrations far below those of biological significance." However, the fact remains that pesticides and pesticide residues are being found in groundwater, and the definition of "biological significance" can change.
In a monitoring study in Suffolk County on Long Island, New York, 13.5 percent of 8,404 wells tested exceeded the state-recommended guideline of 7 ppb aldicarb (Zaki et al., 1982). Of these, 52 percent contained 8 to 30 ppb, 32 percent contained 31 to 75 ppb, and 16 percent contained more than 75 ppb aldicarb. Since then, aldicarb has been found in groundwaters in other areas, particularly in areas with irrigation on sandy soils.
The levers of aldicarb found may not cause health problems, although knowledge on this point is incomplete. Nevertheless, chemical contamination has occurred even though studies were conducted and measures were taken to avoid leaching of chemicals to groundwaters. Once groundwater is contaminated, it may be extremely difficult to decontaminate. For example, there is controversy on how long it will take for aldicarb residues to disappear from groundwater, even after its use is discontinued (Marshall, 1985).
Many people, for various reasons, are apprehensive about situations involving synthetic chemicals as contaminants in the water supply or food chain. It is becoming increasingly evident that these chemicals will be very difficult to keep out of the water or food chain if they are used extensively in crop production. Intensive monitoring and predictive programs are helpful but not foolproof, and these programs are very expansive. With the aldicarb contamination, however, it does appear that improper applications cannot be ruled out as a cause of many such problems.
Both erosion and chemical pollution of soil and water have been accelerated by modem farming practices that do not incorporate conservation tillage practices (King, 1983; Papendick and Elliott, 1984b). The shift in recent decades away from soil-conserving crop rotations to intensive monoculture of grain or continuous row-cropping systems accompanied by intensive tillage has been particularly responsible (Baker, 1985). The availability of pesticides and cheap fertilizer N that originated in the 1950s made these practices feasible and economical for crop production.
Crop rotations that included legumes and grass for hay and pasture became less important on much of the nation's cropland when inexpensive nutrients and pesticides became available. The benefits that soil-conserving rotations generally confer were lost. These included improved soil water infiltration, higher organic master content, increased water-holding capacity, and less erosion. Erosion with continuous row cropping and conventional tillage may be more than double that with sod crops in the rotation. By alternating row crops with cover crops (e.g., hay), erosion in many situations can be held to tolerable levers (Papendick and Elliott, 1984b).
With intensive grain cropping, the use of N fertilizer and pesticides increased dramatically. Nitrogen fertilizer use in the United States increased from 2.4 million mt in 1959 to 11.9 million mt in 1981 (USDA, 1978, 1984) (Figure 1). Use declined after 1981 to about 9.2 million mt in 1983. According to CAST (1985), agriculture used 385,560 mt of pesticides in 1980; the expected use in 1985 is double this amount (Chesters and Schierow, 1985). Herbicide use during the 1970s and into the early 1980s has increased for row and small grain crops but not for forage crops (Table 1). However, the treated acreage of forage crops is very small compared with that for other crops.
The shift to large-scale equipment since the 1960s has also often led to abandonment of erosion control methods such as terracing, grassed waterways, tree shelter belts, and in some cases, even strip cropping. These conservation practices often interfere with the operation of large equipment in large fields (Sampson, 1981). Moreover, making fields larger to accommodate large equipment has reduced border areas that serve to check erosion.
Most recent efforts to control runoff and erosion in large-scale agricultural production systems have focused on conservation tillage, which is defined as any practice that reduces soil or water loss compared with moldboard plowing. It usually consists of direct seeding and fertilizing into the previous crop stubble (no-till) or minimum tillage before seeding, which retains surface residues and either disturbs the seedbed soil only slightly or produces a rough, cloddy seedbed.
1. Erosion and runoff control. Conservation tillage can markedly reduce runoff and erosion over a wide range of climatic, soil, and cropping conditions. Crop residues reduce raindrop impact and, in combination with a rough soil surface, increase infiltration and decrease runoff and sedimentation. In Nebraska studies, for example, no-till systems reduced sheet and rill erosion by 90 percent of that which occurs with conventional moldboard plowing or with other clean-tillage, residue-free systems (Dickey et al., 1984).
Gebhardt et al. (1985) reported that erosion in the Midwest is reduced by 75 percent with surface corn residues and that even with the highly erodible corn (Zea mays)-soybean (Glycine max) rotations, conservation tillage can reduce erosion by at least 50 percent compared with that occurring in conventional systems. In the Palouse wheatlands of eastern Washington and Idaho, no-till (Triticum aestivum) with some stubble can reduce erosion from 45 to 100 mt/ ha to less than 5 mt/ha (USDA, 1978). It seems that conservation tillage is one of the most effective and least expensive methods for controlling soil erosion. When properly applied, it can significantly reduce soil losses.
There are situations in which conservation tillage may be inadequate by itself for acceptable control of erosion. In the Pacific Northwest, erosion from no-till following a pea (Pisum sativum) or lentil (Lens culinaris) crop can be relatively, high, especially on steeper land. These crops produce low amounts of residues and the residues decompose much faster than small grain cereal straw. Hence they lose their effectiveness for controlling erosion much earlier. Secondly these grain legumes tend to produce,a loose, mellow soil surface (somewhat similar to the well-known condition for soybeans), which is highly vulnerable to erosion (Moldenhauer et al., 1983). Erosion control following these crops may require supporting practices such as divided-slope or cross-slope farming and early fall planting of winter wheat in combination with no-till seeding.
Conservation tillage is also often inadequate to control gully or concentrated flow erosion because surface crop residues are not always effective under these conditions. Diversions, terraces grassed waterways, and other structure. are often needed where this is a problem
There are other situations in which conservation tillage alone is insufficient to control erosion, particularly when crop residues are limited, slopes are steep, or runoff occurs over frozen or partly-thawed soils.
2. Chemical pollution. Concerns about conservation tillage are centered on the potential for increased use of pesticides and the undesirable ecological effects that can result. With reduced tillage farmers may rely more on broad-spectrum herbicides such as glyphosate and paraquat for preplant application. Use of other herbicides probably changes less because they are commonly used in conventional tillage systems. Moreover, the chemicals applied in conservation tillage systems are more likely to remain at the soil surface rather than being incorporated into the soil, a situation which raises questions about the escape of these chemicals into the environment via runoff. How conservation tillage affects leaching, volatilization, and runoff of pesticides and fertilizers is not entirely clear.
Many pesticides, particularly herbicides used in conservation tillage, tend to bind strongly to soil particles so that most pollution from treated soil is limited to sediments carried in runoff. It is generally known that N and P are carried in the soil at much higher rates than in runoff water. Baker and Laflen (1983) reported that conservation tillage decreased the loss of agricultural chemicals to surface and groundwaters because sediment concentration was reduced, thereby keeping the applied chemicals from escaping via soil loss. As the water solubility of a surface-applied pesticide increases (when used for no-till), higher concentrations can be expected in runoff than if the chemical is mixed with the soil (Wauchope, 1978). However, solubility alone does not ensure that the pesticide is free to move with runoff water, because there is still a tendency for these chemicals to react with the organic matter in the surface soil.
Much of the uncertainty concerning ,chemical pollution from conservation tillage is associated with subsurface flow and the potential for contaminating groundwater. Overall, conservation tillage decreases evaporation and runoff and increases water infiltration (Baker and Laflen, 1983). Greater infiltration may increase leaching to groundwaters. his may explain the deep movement of (NO3 with reduced tillage that has sometimes been reported. However, this does not satisfactorily explain the downward movement of chemicals that are tightly adsorbed to soil particles, such as some pesticides.
There is a need to learn more about the possible harmful side effects of conservation tillage if pesticides are used more intensively. Few of the ecological effects of pesticide use are well understood, and current research programs are inadequate to provide the necessary answers (Randall, 1984).
Alternative Agriculture (Organic Farming)
A growing interest in alternative farming systems has occurred, especially in organic farming, as a means of addressing the environmental concerns associated with modern large-scale agriculture. In a broad sense, "alternative" and "organic" refer to a spectrum of low-chemical, resource- and energy conserving farming systems and technologies (Youngberg et al., 1984). Organic farming has been defined as "a production system which avoids or largely excludes the use of synthetically compounded fertilizers, pesticides, growth regulators, and livestock feed additives. To the maximum extent feasible, organic farming systems rely upon crop rotations, crop residues, animal manures, legumes, green manures, off-farm organic wastes, mechanical cultivation, mineral-bearing rocks, and aspects of biological pest control to maintain soil productivity and tilth, to supply plant nutrients, and to control insects, weeds, and other pests" (USDA, 1980). Youngberg et al. (1984) stated that although the terms "organic" and "alternative" are used more or less synonymously, "alternative" encompasses terms such as "biological , " " ecological , " " regenerative," and "natural." "Organic" seems to come closest to being a generic term for agricultural systems that place low dependence on the use of manufactured fertilizers and other chemicals for crop and livestock production.
Organic farming differs considerably from conventional farming, mainly in tillage methods, crop rotations, and in the way that crop nutrients are supplied and pests are controlled. Organic farming has potential for reducing some of the negative impacts of conventional agriculture on the environment.
1. Erosion and runoff control. According to a USDA (1980) report, many of the organic farms studied in the United States used soil and crop management practices that are "best management practices" for controlling soil erosion and water pollution. These included sod-based rotations, cover crops, green manure crops, contour farming, and tillage methods that conserve surface residues. Many organic farmers maintain from 25 to 40 percent of their cropland in sod crops, such as grasses and legumes, and have proportionately less cropland in row crops than conventional farmers. With this proportion of sod crops in the rotation, the average annual soil loss is only one-third to one-eighth that which occurs with conventional tillage and continuous row cropping. Even with no-till planting, soil erosion from a corn-soybean sequence tends to be greater than with sod crops in rotations using conventional tillage (USDA, 1980). No-till continuous corn has an average long-term erosion rate comparable to that of a meadow rotation using conventional tillage (USDA, 1980).
Sod or forage legume-based rotations reduce soil loss in two ways. First, soil loss from meadow is negligible after establishment because of the dense vegetative cover and rooting. Second, when the sod or legume is broken out, the residual effects maintain infiltration and leave the soil less erodible (Stewart et al., 1976). The residual effect of meadow may continue, though on a declining level, for two to three years and carry through the entire small-grain and rowcropping rotations typically used by organic farmers.
Cover crops, which are also routinely used by organic farmers, can reduce soil loss nearly 50 percent when they follow crops that leave little residue cover after harvest: e.g., corn harvested for silage, some grain legumes, and most vegetable crops. For example, the potential erosion for continuous conventional tilled corn with stover removed was about 20 percent less where an overwinter cover crop was used compared with situations involving no cover crops (Stewart et al., 1975).
Most organic farmers employ shallow tillage (no more than 6 to 10 cm deep) to mix the crop residues, manures, or other organic wastes in the upper soil layers. This is generally accomplished with chisel plow or disk-type implements and is followed by one or more secondary operations with a light disk or field cultivator. This noninversion tillage is highly effective in controlling erosion and reducing sedimentation and nutrient runoff.
The use by many organic farmers of meadow rotations, repeated applications of animal manures or other organic wastes, and incorporation of green manures into soil tends to conserve or increase soil organic matter. This in turn may decrease soil erosion through improved soil physical and biological properties. Bolton et al. (1985) showed that soil biological activity and organic matter were higher in soil from a farm on which mainly organic agriculture techniques were used as compared with soil from an adjacent conventional farm. According to Wischmeier and Smith (1978), an increase in the soil organic matter content of one percent decreases the erosion potential by about 10 percent.
2. Chemical pollution. Organic farmers rely almost entirely on nonchemical methods for nutrient supply and crop protection. The main source of N is from legumes in the rotation. Unprocessed insoluble minerals such as greensand and phosphate rock are used when needed to supply K and P (USDA, 1980). Much of the grain and forage on organic farms is used for animals and the crop residues and manures are returned to the land, reducing the need for chemical fertilizer. Nitrogen leaching is quite unlikely for an organic farm except where manure has been heavily applied or a legume such as alfalfa has been grown for a long period and is broken out and intensively tilled.
The diversity of crops in the rotation is a means of controlling weeds, insects, and diseases. Weeds are further controlled with tillage, mowing, and grazing. Biological methods of insect control with organisms such as Bacillus thuringiensis and ladybird beetles (Coccinellidea ) are employed occasionally (Hoy and Herzog, 1985). Since organic systems tend to avoid or largely exclude fertilizers and pesticides, they do not contribute such chemicals to environmental pollution.
Organic farmers are usually not advocates of no-till because pesticides are required with present technology. They are also concerned that leaving too much crop residue on the surface will increase pest problems. Organic farmers question the long-term sustainability of any agricultural system that depends heavily on pesticide use. Their questions deal primarily with cost, effectiveness, and stability of the ecosystem. Moreover, many farmers are convinced that some tillage is necessary to maintain soil aeration, to stimulate microbial activity and organic matter decomposition, and to decrease compaction.
During early settlement of the United States, wild animals were an important source of meat for the pioneers. Two major industries, logging and agriculture, drastically altered the environment
for wildlife. Some species that could not tolerate human activity or were overhunted were considerably reduced in numbers, examples being the bison (Bison bison) and elk (Cervus elaphus). Other species such as the fox squirrel (Sciurus niger), the bobwhite (Colinus virginianus), and the prairie chicken (Tympanuchus cupido) expanded their range and increased in numbers with early agriculture. Thus, it seems that food originally was a limiting factor controlling the numbers of some species. Early agriculture actually contributed to diversification of the habitat. As agriculture became more pervasive, natural habitat was extensively converted to crop fields and wildlife populations decreased. This was especially true for the prairie chicken, which has disappeared from much of its former range.
The ring-necked pheasant (Phasianus colchicus) was imported from Asia to Oregon in 1881 (Lauckhart and McKean, 1956). From this successful population in the Northwest, birds were trapped and moved to establish populations elsewhere in the northern part of the United States. In this country, farming also was friendly to the pheasant because it provided the weeds, "edge effect," and waste grains that constitute a favorable habitat for this species.
Changes in farming in the United States have caused many wildlife species to decline in numbers. Each species has distinct habitat requirements for its life cycle and because food is one of those requirements, farming has benefited some wildlife species. However, intensive farming is not compatible with wildlife when row crops replace former marshes, grasslands, and forests necessary for wildlife to breed, nest, and be protected from disturbance and predators.
Specialized and sophisticated farm machinery, advances in plant breeding, and the development of rapid transportation have contributed to regionalization of food crops. All these factors have reduced local habitat diversity within the agricultural system. Wildlife has consequently declined in both numbers and richness (numbers of species).
Pheasants were successfully introduced to the Midwest from about the turn of the century until the 1920s, and their numbers increased greatly. Within a few years after introduction, the first hunting seasons were held. Farming the Midwest until the fate 1940s and 1950s had many characteristics of organic farming. Farms were diversified and a variety of crops was grown. Livestock were raised on nearly every farm and horses were used extensively until the 1930s and into the 1940s.
Farms typically had fencerows weeds, headlands (a place at the end of the field for machinery to turn), undrained marches, and cattle and horses pasturing waste crops and breaking the crust of snow for foraging birds. On such farms, pheasants could easily find food a secure place to incubate eggs, and cover to shelter them from storms and predators.
Pheasants responded to these conditions with the development of huge populations. In South Dakota in the 193os and 1940s, which then claimed to be the "Pheasant Capital of the World," pheasants were hunted for as long as 120 days with a bag limit of 10 birds per day. (Hipschman, 1959). Probably there were at least 50 million birds in the state in the 1940s, but with changes in agriculture, that figure decreased to 2 million in 1966 (Dahlgren, 1967).
The decline in pheasant population occurred for several reasons: mars drainage, development of farm machinery that led to removal of fencerows and headlands, removal of livestock an barns from many farms, herbicide modification of vegetation, large fields, and monotypes of cultivated row crops. The detail of these changes in land use was described by Mohlis (1974) for north central Iowa and by Fischer (1974) for a small research area in northern Iowa The latter 615-ha area, the Winnebago Research Area, had a pheasant population in the fall of 1941 estimated to be as high as 950 (Baskett, 1947), but only 28 in 1973 (Fischer, 1974). Farris et al (1977) stated that while 59 percent o the land on the Winnebago Research Area had potential nesting, cover during 1939-41, the figure had decreased to only 14 percent in 1973 (nearly 45 percent of which was in an annual USDA set-aside program). Changes in habitat and pheasant populations have been detailed for Nebraska by Taylor et al. (1978) and for prairie chickens, bobwhites, and cottontail (Sylvilagus floridanus) in Illinois by Vance (1976).
Clearly, farming in America has changed the face of the landscape. Not all wildlife species have been adversely affected (e.g., white-tailed deer (Odocoileus virginianus), but the decline of some species dependent on the farm illustrates how drastically agriculture has changed.
Biologists have been concerned about pesticides and their effects on wildlife for several reasons. For example, whereas farm workers can be told not to enter a freshly sprayed field, wildlife species cannot be given the same warning. Application of organochlorines has resulted in behavioral changes in pheasant chicks (Dahlgren et al., 1970, 1972) that were mediated through both cocks and hens. When either parent was administered dieldrin or polychlorinated biphenyls (PCBs, an industrial chemical), offspring did more poorly than offspring of untreated birds on a visual-cliff test (used to test whether chicks would jump to the visually close or visually deep side). The effect was greatest where both the cock and hen received the chemical. Similarly, offspring of treated parents were more easily captured by hand. In the wild, birds must depend on innate behavior patterns for survival, such as the ability to avoid capture by predators. Use of DDT, dieldrin, and PCBs has been effectively controlled in the United States, but not before wildlife populations were exposed to possible behavioral effects and the documented eggshell thinning that endangered predatory birds and brown pelicans (Pelecanus occidentalis) (Klaas, 1982). Because chemicals were developed and in common use before their danger was recognized, and because full evaluation of any hazards they might pose is so costly that it might not be done, most wildlife biologists would rather do without chemicals. However, many herbicides and some short-lived insecticides seem to be relatively safe (Klaas, 1982). The concern of some wildlife biologists with herbicides is that they destroy the vegetative diversity, especially that contributed by broad-leafed plants.
Wildlife exists in a great variety of farm habitats: the Palouse of the Northwest, irrigated fields of the West, wheat fields in Kansas, corn and soybean fields in Iowa, and rice fields in Louisiana. However, wildlife could be benefited by a shift toward diversity of crops grown on farms or by a real effort to reduce soil erosion. Major economic or political forces, such as a large increase in transportation costs or a societal decision that we will not continue to tolerate high soil erosion rates, are probably the only ones potent enough to cause such changes.
Conservation tillage is more beneficial for controlling soil erosion than for enhancing wildlife (Cacek, 1984; Wooley et al., 1985). The likely increases in chemical use and the disturbance of the field during the nesting season are chief obstacles to a wildlife benefit from conservation tillage. However, no-till is an exception. If farmers enter the field only once early in the reproductive season for planting, fertilization, and weed control, wild birds can reproduce successfully. Wildlife can benefit from improved weed control by undercutting in wheat field stubble without destroying all bird nests (Rodgers and Wooley, 1983).
A trend toward organic farming would produce diversity in crop types and smaller fields, benefiting many nongame and game birds. As an example, we can contras" two nesting studies made in divergent farm habitats--one by Trautman (1960) in eastern South Dakota, and the other by Linder et al. (1960) in south-central Nebraska. The habitat in South Dakota was more diverse than the simpler habitat (chiefly wheat and row crops) in Nebraska (Table 2). The South Dakota habitat supported a spring population of 110 hens/ section (2.59 km2) compared with 21 hens/section in Nebraska. Nesting cover, as winter wheat, pasture, and alfalfa (Medicago saliva), comprised 37 percent of the total land area and produced about 63 percent of all pheasant chicks produced in the simpler habitat of Nebraska. Nesting cover, as oats (Avena saliva), pasture, alfalfa, barley (Hordeum vulgare), and wheat constituted about 48 percent of the land area in South Dakota and produced about 54 percent of all pheasants. Diversity of habitat and smaller fields are more favorable to pheasant production and were associated with higher populations.
Alternative farming systems protect the environment and benefit wildlife. The extensive crop rotations, which include legumes for forage and green manure, cover crops, and meadows, increase soil resistance to erosion and soil organic master content, when compared with conventional intensive tillage, monoculture, and row crop farming methods. Alternative farming practices, in most cases, will reduce soil loss below the soil loss tolerance (T) value. The reduced or non-use of manufactured chemicals with these farming systems greatly decreases the environmental hazards and possible adverse effects on wildlife. Wildlife generally benefits from these farming systems because crops are more diverse and offer a mixture of habitats.
Acknowledgments: R. I. Papendick and L. F. Elliott, USDA-ARS, in cooperation with College of Agriculture and Home Economics, Washington State University, Pullman, WA 99164. Scientific paper No. 7360. The research of R. B. Dahlgren was supported jointly by the U.S. Fish and Wildlife Service, Iowa Conservation Commission, Iowa State University, and Wildlife Management Institute. (Journal paper. J12133 of the Iowa Agriculture and Home Economics Experiment Station, Ames, Project No. 2412).
1. Baker, D. B. 1985. Regional water-quality impacts of intensive row-crop agriculture: A Lake Erie Basin case study. J. Soil Water Conserv. 40:125-132.
2. Baker, J. L., and J. M. Lafnen. 1983. Water quality consequences of conservation tillage. J. Soil Water Conserv. 38:186-193.
3. Baskett, T. S. 1947. Nesting and production of the ring-necked pheasant in north-central lowa. Ecol. Monogr. 17: 1-30.
4. Bolton, H., Jr., L. F. Elliott, R. I. Papendick, and D. F. Bezdicek. 1985. Soil microbial biomass and selected soil enzyme activities: Effect of fertilization and cropping practices. Soil Biol. Biochem. 17: 297302.
5. Brown, L. R., and E. C. Wolf 1984. Soil erosion: Quiet crisis in the world economy. Worldwatch Paper 60. Worldwatch Institute, Washington, D.C. 49 pp.
6. Cacek, T. 1984. Organic farming: The other conservation farming system. J. Soil Water Conserv. 39:357360.
7. CAST (Council for Agricultural Science and Technology). 1985. Agriculture and groundwater quality. CAST Report No. 103, pp. 6, 8, 10, II, 13. Iowa State University Station, Ames. 6b pp.
8. Chesters, G., and L. Schierow. 1985. A primer on nonpoint pollution. J. Soil Water Conserv. 40:9-13.
9. Clark, E. H., II. 1985. The off-site costs of soil erosion. J. Soil Water Conserv. 40:19-22.
10. Cohen, S. V., S. M. Creeger, R. F. Carsel, and C. F. Enfield. 1984. Potential pesticide contamination of groundwater from agricultural uses. In: Treatment and Disposal of Pesticide Wastes, R. F. Krueger and J. N. Seiber (eds.), pp. 297-325. Am. Chem. Soc. Symp. Ser. 259.
II. Crosson, P. R., and A. T. Stout. 1983. Productivity effects of cropland erosion in the United States. Resources for the Future, Washington, D.C. 103 pp.
12. Dahlgren, R. B. 1967. The pheasant decline. South Dakota Department of Game, Fish and Parks, Pierre. 44 pp.
13. Dahlgren, R. B., R. L. Linder, and C. W. Carlson. 1972. Polychlorinated biphenyls: Their effects on penned pheasants. Environ. Health Perspect., Exp. Issue No. 1:89-101.
14. Dahlgren, R. B., R. L. Linder, and K. K. Ortman. 1970. Dieldrin effects on susceptibility of penned pheasants to hand capture. J. Wildl. Manage. 34:957 959.
15. Dickey, E. C., D. P. Shelton, and T. R. Peterson. 1984. Conservation tillage: Effective, inexpensive erosion control. Farm, Ranch, and Home Quarterly 30:18-20. University of Nebraska, Lincoln.
16. Duda, A. M., and R. J. Johnson. 1985. Cost-effective targeting of agricultural nonpoint-source pollution controls. J. Soil Water Conserv. 40:108-11.
17. Farris, A. L., E. D. Klonglan, and R. C. Nomsen. 1977. The ring-necked pheasant in lowa. Iowa Conservation Commission, Des Moines. 147 pp.
18. Fischer, W. A. 1974. Nesting and production of the ring-necked pheasant on the Winnebago Research Area, lowa. M.S. thesis. Iowa State University Library, Ames. 63 pp.
19. Foster, G. R., R. A. Young, M. J. M. Romkens, and C. A. Onstad. 1985. Processes of soil erosion by water. In: Soil Erosion and Crop Productivity, R. F. Follett and B. A. Steward (eds.), pp. 137-162. American Society of Agronomy, Madison, Wl.
20. Frye, W. W., O. L. Bennett, and G. J. Buntley. 1985. Restoration of crop productivity on eroded or de graded soils. In: Soil Erosion and Crop Productivity R. F. Follett and B. A. Stewart (eds.), pp. 335-356 American Society of Agronomy, Madison, Wl.
21. Gebhardt, M. R., T. C. Daniel, E. E. Schweizer, and R. R. Allmaras. 1985. Conservation tillage. Science 230:625-630.
22. Hallberg, G. 1984. Agricultural chemicals and groundwater quality in lowa. Coop. Ext. Serv. Rpt. CE-2081j, lowa State University, Ames. 6 pp.
23. Hauck, R. D., and M. Koshino. 1971. Slow-release and amended fertilizers. In: Fertilizer Technology and Use, 2nd ed., R. A. Olson, T. J. Army, J. J. Hanway, and V. J. Kilmer (eds.). Soil Science Society of America, Madison, Wl.
24. Hipschman, D. 1959. A look at the past half century. South Dakota Conserv. Digest. 26(2):20-31,37.
25. Hoy, M. A., and D. C. Herzog (eds.). 1985. Biological control in agricultural IPM systems. Academic Press, New York. 589 pp.
26. King, A. D. 1983. Progress in no-till. J. Soil Water Conserv. 38:160-161.
27. Klaas, E. E. 1982. Effects of pesticides on non-target organisms. In: Proceedings of the Midwest Agricultural Interfaces with Fish and Wildlife Resources Workshop, R. B. Dahlgren (compiler), pp. 7-9. Iowa Coop. Wildl. Research Unit, lowa State University, Ames.
28. Langdale, G. W., R. A. Leonard, and A. W. Thomas. 1985. Conservation practice effects on phosphorus losses from Southern Piedmont watersheds. J. Soil Water Conserv. 40:157-161.
29. Larson, W. E. 1981. Protecting the soil resource base. J. Soil Water Conserv. 36:13-16.
30. Lauckhart, J. B., and J. W. McKean. 1956. Chinese pheasants in the Northwest. In: Pheasants in North America, D. L. Allen (ed.), pp. 43-89. American Wildlife Institute, Washington, D.C.
31. Linder, R. L., D. L. Lyon, and C. P. Agee. 1960. An analysis of pheasant nesting in south-central Nebraska. Trans. North Am. Wild. Conf 25:214-230.
32. Marshall, Eliot. 1985. The rise and decline of Temik. Science 229:1369-1371.
33. Mohlis, C. K. 1974. Land use and pheasant habitat in north-central lowa, 1938-1973. M.S. Thesis. Iowa State University Library, Ames. 84 pp.
34. Moldenhauer, W. C., G. W. Langdale, W. Frye, D. K. McCool, R. 1. Papendick, D. E. Smika, and D. W. Fryrear. 1983. Conservation tillage for erosion control. J. Soil water Conserv. 38:144-151.
35. Myers, C. F., J. Meek, S. Tuller, and A. Weinberg. 1985. Nonpoint sources of water pollution. J. Soil Water Conserv. 40:14-22.
36. Papendick, R. 1., and L. F. Elliott. 1984a. Soil physical factors that affect plant health. In: Challenging Problems in Plant Health, Thor Kommedahl and P. H. Williams (eds.), pp. 168-180. The American Phytopathological Society, St. Paul, MN.
37. Papendick, R. 1., and L. F. Elliott. 1984b. Tillage and cropping systems for erosion control and efficient 60 nutrient utilization. In: Organic Farming: Current Technology and Its Role in a Sustainable Agriculture, D. F. Bezdicek and J. F. Power (eds.), pp. 69-81. American Society of Agronomy, Madison, Wl.
38. Piest, R. F., R. G. Spomer, and P. R. Muhs. 1977. A profile of soil movement on a cornfield. In: Soil Erosion: Prediction and Control. pp. 160-166. Spec. Publ. No. 21. Soil Conservation Society of America, Ankeny, IA.
39. Power, J. F. 1983. Research in agricultural ecosystems, North Central United States. In: Nutrient Cycling in Agricultural Ecosystems, R. Lowrance, R. L. Todd, L. Ausmussen, and R. Leonard (eds.). p~ 121-133. Coll. of Agric. Spec. Publ. 33, University c Georgia Athens.
40. Randall G. W. 1984. Factors limiting the efficacy of agricultural chemicals in changing agricultural production systems--Current state of the art. In: Changing Agricultural Production Systems and the Fate of Agricultural Chemicals, G. W Irving, Jr (ed. ). pp. 76-87. Agricultural Research Institute, Bethesda MD.
41. Rodgers, R. D., and J. B. Wooley. 1983. Conservation tillage impacts on wildlife. J. Soil Water Conserv 38:212-213.
42. Sampson, R. N. 1981. Farmland or wasteland. Rodale Press, Emmaus, PA. 422 pp.
43. Schepers, J. S., and L. N. Mielke. 1983. Nitrogen fertilization, mineralization, and leaching under irrigation in the Midwest. In: Nutrient Cycling in Agricultural Ecosystems, R. Lowrance, R. Todd, L Asmussen, and R. Leonard (eds.). pp. 325-334. Georgia Agric. Exp. Stn. Spec. Publ. 23. University o Georgia, Athens.
44. Schuman, G. E., T. M. McCalla, K. E. Saxton, and H. T. Knox. 1975. Nitrate movement and its distribution in the soil profile of differentially fertilizer corn watersheds. Soil Sci. Soc. Am. Proc. 39:1192 1197.
45. Stewart, B. A., D. A. Woolhiser, W. H. Wischmeier J. H. Caro, and M. H. Frere. 1975. Control of water pollution from cropland, Vol. I. U.S. Department of Agriculture and Environmental Protection Agency, Washington, D.C. 111 pp.
46. Stewart, B. A., D. A. Woolhiser, W. H. Wischmeier J. H. Caro, and M. H. Frere. 1976. Control of water pollution from cropland, Vol. II. U.S. Department of Agriculture and Environmental Protection Agency, Washington, D.C. 187 pp.
47. Stoltenberg, N. L., and J. L. White. 1953. Selective losses of plant nutrients by erosion. Soil Sci. Soc. Am. Proc. 17:406-410.
48. Taylor, M. W., C. W. Wolfe, and W. L. Baxter. 1978. Land-use change and ring-necked pheasants in Nebraska. Wildl. Soc. Bull. 6(4):226-230.
49. Trautman, C. G. 1960. Evaluation of pheasant nesting habitat in eastern South Dakota. Trans. North Am. Wildl. Conf 25:202-213.
50. U.S. Department of Agriculture. 1978. Agricultural statistics. U. S. Government Printing Office, Washington, D.C.
51. U.S. Department of Agriculture. 1980. Report and recommendations on organic farming. A special report prepared for the Secretary of Agriculture. U. S. Government Printing Office, Washington, D.C. 164 PP.
52. U.S. Department of Agriculture. 1981. Soil, water and related resources in the United States: Status, conditions, and trends: 1980 RCA Appraisal, Part 1. U. S. Government Printing Office, Washington, D.C.
53. U.S. Department of Agriculture. 1984. Agricultural statistics. U. S. Government Printing Office, Washington, D.C.
54. U.S. Environmental Protection Agency. 1976. Quality criteria for water. Washington, D.C.
55. U.S. Environmental Protection Agency. 1983. Chesapeake Bay program: Findings and recommendations. Region 3, Philadelphia PA.
56. Vance, D. R. 1976. Changes ;n land use and wildlife populations in southeastern Illinois. Wildl. Soc. Bull. 4(1):1 1-15.
57. Wauchope, R. D. 1978. The pesticide content of surface water draining from agricultural fields--A review. J. Environ. Qual. 7:459-472.
58. Wischmeier, W. H., and D. D. Smith. 1978. Predicting rainfall erosion losses--A guide to conservation planning. U. S. Dept. Agric. Handb. No. 537. U. S. Government Printing Office, Washington, D. C. 164 pp.
59. Wooley, J. B., Jr., L. B. Best, and W. R. Clark. 1985. Impacts of no-till row cropping on upland wildlife. Trans. North Am. Wildl. Natur. Resourc. Conf 50:157-168.
60. Youngberg, 1. G., J. F. Parr, and R. 1. Papendick. 1984. Potential benefits of organic farming practices for wildlife and naturel resources. Trans. North Am. Wildl. Nat. Resour. Conf 49:141-153.
61. Zaki, M. H., D. Moran, and D. Harris. 1982. Pesticides in groundwater: The aldicarb story in Suffolk County, New York. Am. J. Public Health 72:13911 395.
Citation : Papendick, Robert I., Elliot Lloyd F., Dahlgren, Robert B., 1986, " Environmental consequences of modern production agriculture : how can alternative agriculture address these issues and concerns ?", Vol. 1, No. 1, pp. 3-10.
Copyright © 1986
Reprinted with permission.
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